This document summarizes a literature review on the reproductive and developmental toxicity of phthalates towards fish. It discusses common phthalates like DEHP, DBP, DINP and DIDP. The review finds that DEHP and DBP can cause effects like decreased gonad size and vitellogenin production in fish. More recent studies show DINP and DIDP may also have endocrine disrupting effects. The document concludes some phthalates pose risks to fish reproduction and more research is needed, especially on emerging phthalates.

1.
A
review
of
phthalates
and
the
associated
reproductive
and
developmental
toxicity
towards
fish.
Masters
literature
thesis
-­‐
12
EC
Emma
Greenwell
(10407995)
Biological
sciences:
Limnology
and
oceanography
Supervisor:
Liana
Bastos
Sales
Examiner:
Michiel
Kraak
20th
December
2013
–
27th
March
2014

2. 2
Table
of
Contents
I.
Abstract
..............................................................................................................
4
II.
Introduction
.......................................................................................................
5
2.1
What
are
phthalates?
................................................................................................................................
5
2.1.1
Common
phthalates
...............................................................................................................................
5
2.2
Environmental
fate
of
phthalates
........................................................................................................
6
2.2.1
Differences
in
seasons
............................................................................................................................
8
2.3
Levels
in
the
environment
......................................................................................................................
8
2.4
Half-­‐lives
.........................................................................................................................................................
9
2.5
Inside
the
organism
................................................................................................................................
11
2.6
Modes
of
action
once
inside
an
organism
.....................................................................................
11
2.7
Environmental
risk
limits
....................................................................................................................
12
2.8
Objective
......................................................................................................................................................
12
III.
Method
...........................................................................................................
13
IV.
Results
............................................................................................................
13
4.1
Summary
of
literature
(1980-­‐1999)
...............................................................................................
14
4.2
Literature
(1980-­‐1999)
........................................................................................................................
16
4.3
Summary
of
literature
post
2000
.....................................................................................................
17
4.4
Literature
post
2000
..............................................................................................................................
19
4.4.1
DEHP
..........................................................................................................................................................
19
4.4.2
DBP
.............................................................................................................................................................
24
4.4.3
DEHP
and
DBP
.......................................................................................................................................
28
4.4.4
DINP
and
DIDP
......................................................................................................................................
29
V.
Discussion
........................................................................................................
29
5.2
DEHP
.............................................................................................................................................................
30
5.3
DBP
................................................................................................................................................................
30
5.4
Nominal
concentration
experiments
with
DEHP
and
DBP
....................................................
31
5.5
DINP
and
DIDP
..........................................................................................................................................
32
5.6
Exposure
routes
.......................................................................................................................................
33
5.7
Problematic
variables
and
environmental
risk
limits
.............................................................
33
VI.
Conclusions
.....................................................................................................
34
6.1
Classification
of
phthalates
..................................................................................................................
34
6.1.1
DEHP
..........................................................................................................................................................
35
6.1.2
DBP
.............................................................................................................................................................
35
6.1.3
DINP
and
DIDNP
...................................................................................................................................
35
4.9
Recommendations
...................................................................................................................................
35
VII.
Author’s
remarks
...........................................................................................
36
VIII.
References
....................................................................................................
37

3. 3
GLOSSARY
Environmental
risk
limit
(ERL)
–
represent
the
potential
risk
of
the
substance
to
the
ecosystem
and
are
derived
using
data
from
ecotoxicology
and
environmental
chemistry.
Oocytes
–
a
cell
in
an
ovary,
which
might
undergo
meiotic
division
to
form
an
ovum.
Vitellogenin
–
a
protein
present
in
the
blood
from
which
the
substance
of
the
egg
yolk
is
derived.
Planktivores
–
An
organism
that
feeds
on
plankton.
Glucuronides
–
any
substance
produced
by
linking
a
glucuronic
acid
to
another
substance
(via
glycosidic
bonds).
This
method
(glucorinidation)
is
used
by
animals
to
help
excrete
toxic
substances
from
the
body.
Environmental
risk
assessmen
(ERA)
–
An
evaluation
of
the
interactions
of
agents,
human
and
ecological
resources.
No
observed
effect
concentration
(NOEC)
–
the
highest
treatment
(test
concentration)
of
a
substance
that
shows
no
statistical
effect
compared
to
a
control.
Predicted
no
effect
concentration
(PNEC)
–
the
concentration
below
which
a
specified
percentage
of
species
in
an
ecosystem
are
expected
to
be
protected.
Nominal
concentration
–
The
concentration
if
you
all
test
material
added
to
the
test
solution
dissolved.
Effective
concentrations
(EC50)
–
the
concentration
of
a
substance,
which
induces
a
response
halfway
between
the
baseline
and
maximum
after
a
specified
exposure
time.
The
number
refers
to
the
position
within
the
baseline-­‐maximum
scale.
Gonado-­‐somatic
index
–
calculation
of
the
gonad
mass
as
a
proportion
of
the
total
body
mass.
Spermatozoa
–
a
sperm
cell.
Spermatocyte
–
immature
male
germ
cell
which
undergoes
meiosis
developme
into
a
sperm
cell.
Spermatagonia
–
any
cell
of
the
male
gonad
that
mature
to
form
spermatocytes.
Hypertrophy
–
a
non-­‐tumorous
enlargement
of
an
organ
(or
part)
as
a
result
of
increased
cell
size
rather
than
cell
number.
Spiggin
–
a
glycoprotein
glue
used
by
three-­‐spined
sticklbacks
to
stick
their
nests
together.
Peroxidation
–
a
chemical
reaction
in
which
oxygen
atoms
are
formed
leading
to
production
of
peroxides.
Photodegradtion
/photodegradable
–
substances
capable
of
being
chemically
broken
down
by
prolonged
exposure
to
light.
Octanol-­‐water
partition
coefficient
(Kow)
–
a
coefficient
representing
the
ratio
of
the
solubility
of
a
compound
in
octanol
to
its
solubility
in
water.
Soil
organic
carbon-­‐water
partitioning
coefficient
(Koc)
–
the
ratio
of
the
mass
of
a
chemical
that
is
adsorbed
in
the
soil
per
unit
mass
of
organic
carbon
in
the
sol
per
the
equilibrium
chemical
concentration
in
solution.
Phytoremediation
–
the
use
of
plants
to
remove/neutralize
contaminants.

4. 4
I.
Abstract
Phthalates
are
endocrine
disrupting
compounds
produced
on
a
mass
scale
for
use
in
plastics.
They
are
not
chemically
bound
to
the
product
and
therefore
leach
into
the
environment
exposing
fish
to
a
range
of
endocrine
toxicities.
Environmental
risk
limits
(ERLs)
are
difficult
to
calculate
as
different
solubility,
exposure
method,
fish
species
and
even
age
all
combine
to
produce
different
toxicity
effects.
In
most
literature
environmental
phthalate
levels
were
above
the
ERL.
This
paper
focuses
on
what
are
associated
endocrine
toxicity
effects
(metabolic,
developmental
and
reproductive)
of
di-­‐2-­‐ethyl-­‐
hexyl
phthalate
(DEHP),
di-­‐butyl
phthalate
(DBP),
di-­‐isononyl
phthalate
(DINP)
and
di-­‐isodecyl
phthalate
(DIDP).
Results
consist
of
18
studies
on
phthalate
toxicity
filtered
to
only
include
results
from
DEHP,
DBP,
DINP
and
DIDP
on
fish
species.
A
mixture
of
effects
on
growth
inhibition,
VTG
level
alteration,
inhibition
of
oocyte
maturation,
increased
mortality,
spinal
deformities
and
maturation
inducing
hormone
alterations
etc.
were
observed
with
all
both
DEHP
and
DBP.
Effects
were
seen
to
be
more
potent
in
pre/early
life
exposure
compared
to
adults
and
sometimes
even
irreversible.
Both
DEHP
and
DBP
phthalates
produces
developmental
toxicity
effects
such
as
increased
mortality,
retardation
in
ovary
development,
decreases
in
body
weight
and
length,
inhibition
of
5α-­‐adione,
decreases
in
fertility
and
many
more.
The
order
of
literature
available
went
DEHP>DBP>DINP/DIDP.
For
the
latter
two
(DINP/DIDP)
only
one
study
was
found
post
year
2000.
The
availability
of
DEHP
and
DBP
information
allows
to
derive
reasonable
ERLs
values.
However
due
to
the
lack
of
DINP/DIDP
information
DEHP
is
used
as
a
proxy
for
DINP/DIDP
ERLs.
In
conjunction,
there
is
no
uniform
exposure
route
to
which
ERL’s
are
based
on
and
as
seen
in
the
results
different
exposure
routes
of
the
same
compound
can
produce
different
effects.
More
solid
guidelines
of
phthalate
testing
are
needed
on
all
compounds
especially
those
of
DINP
and
DIDP.

5. 5
II.
Introduction
2.1
What
are
phthalates?
Phthalates
are
chemical
compounds
used
to
reduce
the
chemical
affinity
between
plastic
molecules
therewith
increasing
the
flexibility
of
the
product
sometimes
making
up
50%
of
the
finished
plastic
product
(Oehlmann
et
al.,
2009;
OEHHA,
2009).
They
are
also
known
to
be
endocrine
disrupting
compounds
(EDCs)
(Ikele,
2011).
EDCs
may
be
natural
or
synthetic
compounds
that
interfere
with
endocrine
regulated
processes
such
as
growth
and
reproduction
(Crain
et
al.,
2008).
The
international
program
for
chemical
safety
defines
endocrine
disrupters
as
“exogenous
substances
that
alter
function(s)
of
the
endocrine
system
and
consequently
cause
adverse
health
effects
in
an
intact
organism
or
its
progeny
secondary
to
changes
in
the
endocrine
function”
(ECPI,
2009).
Production
of
phthalates
consists
of
around
1
billion
tones
per
year
worldwide.
They
are
present
in
the
medical
environment,
cosmetics,
computers,
children
toys,
food
packaging,
car
products
and
paint
making
them
an
unavoidable
part
of
modern
life
(Mankidy
et
al.,
2013;
OEHHA,
2009;
Guven
and
Coban,
2013
and
Carnevali
et
al.
2010).
Phthalates
are
not
chemically
bound
to
the
plastic
molecules
within
the
product
meaning
they
are
able
to
leach
out
into
the
environment
rendering
these
compounds
unstable
within
their
plastic
counterpart
(Oehlmann,
et
al.,
2009
and
Mankidy
et
al.,
2013).
Consequently
phthalates
are
ubiquitous
the
environmental
and
ecological
concerns
surrounding
them
are
increasing.
2.1.1
Common
phthalates
The
general
structure
of
phthalates
can
be
seen
in
figure
1
(to
the
right)
(R-­‐alkyl
chain).
The
most
common
phthalates
are
di-­‐n-­‐butyl
phthalate
(DBP)
and
di-­‐2-­‐ethyl-­‐
hexyl
phthalate
(DEHP)
(Jarmolowicz
et
al.,
2013;
Huang
et
al.,
2008
and
Uren-­‐Webster
et
al.,
2010).
These
two
specific
phthalates
Figure
1:
General
structure
of
phthalates
(Ogunfowokan
et
al.,
2006)

6. 6
occur
at
higher
concentrations
than
other
phthalates
(Van
Wezel
et
al.,
2000)
and
have
the
highest
toxicity
(out
8
common
phthalates
under
the
U.S.
environmental
protection
agency
(EPA)
management
plan)
to
terrestrial
and
aquatic
organisms
(EPA,
2012).
These
two
phthalates
produce
reproductive
and
developmental
toxicity
effects
(Jarmolowicz
et
al.,
2013;
Lee
and
Liang
2011
and
Zanotelli
et
al.,
2009).
Newer
phthalate
compounds
such
as
di-­‐isononyl
phthalate
(DINP)
and
di-­‐isodecyl
phthalate
(DIDP)
have
shown
to
have
no
(or
very
low)
toxic
effects
on
aquatic
organisms
(EPA,
2012;
Oehlmann
et
al.,
2009
and
Hallmark
2010)
despite
the
reproductive
development
effects
in
two
generations
of
rats
(OEHHA,
2010).
2.2
Environmental
fate
of
phthalates
Once
in
the
environment
phthalates
are
transported
through
water
where
they
may
be
dissolved
(water
sink)
or
due
to
its
low
solubility
end
up
within
the
sediment
(Huang
et
al.,
2008).
Here
the
phthalate
compounds
are
transferred
to
fish
and
other
aquatic
organisms
through
their
diet
or
by
water
(Jarmolowicz
et
al.,
2009).
Benthic
feeders
contain
higher
levels
of
phthalate
compounds
within
their
system
compared
to
planktivores
due
to
the
low
solubility
of
most
phthalates
(Huang
et
al.,
2008;
Oehlmann
et
al.,
2009;
Mankidy
et
al.,
2013
and
OEHHA,
2009).
The
levels
of
phthalates
within
water
are
affected
by
water
quality
such
as
chemical
oxygen
demand,
dissolved
oxygen,
ammonia-­‐nitrate,
suspended
solids
etc.
(Haung
et
al.,
2008).
Each
phthalate
has
a
different
molecular
weight
that
also
gives
it
different
properties.
A
high
molecular
weight
(HMW)
means
that
the
compound
may
be
less
biologically
available
while
low
molecular
weight
(LMW)
compounds
are
more
biologically
available
(Berge
et
al.,
2013).
This
makes
sense
with
some
literature
as
DBP
(MW
278.4g/mol)
has
a
lower
molecular
weight
then
DEHP
(390.6g/mol)
so
therefore
is
more
available
for
uptake
(Teil
et
al.,
2012).
In
France
three
fish
species
were
analyzed
to
see
which
phthalates
were
more
abundant
(Teil
et
al.,
2012).
Contradictory
to
Huang
et
al.,
2008)
DBP
was
the
main
phthalate
found
in
roach
(Rutilus
rutilus)
followed
by

7. 7
DEHP.
This
would
confirm
the
theory
that
LMW
compounds
are
more
readily
biologically
available
than
HMW.
The
gradients
for
soil
was
however
opposite
with
DEHP
being
the
main
phthalate,
but
this
too
would
fit
theory
that
phthalates
with
a
low
log
Kow
(inverse
of
octanol-­‐water
partition
coefficient,
related
to
aqueous
solubility)
are
better
at
forming
solutes
(dissolving)
than
phthalates
with
a
high
log
Kow.
DBP
has
a
log
Kow
of
4.75
while
DEHP
has
a
higher
one
at
7.5.
Phthalates
with
a
high
log
Kow
are
more
likely
to
have
a
higher
%
in
the
sediment
as
the
particles
that
do
not
dissolve
sink
towards
the
sediment
within
a
water
column
(Berge
et
al.,
2013).
As
DEHP
has
a
higher
log
Kow
it
means
that
it
will
be
present
in
larger
quantities
compared
to
DBP
in
sediment
samples.
When
looking
at
the
log
Kow
of
DINP
and
DIDP
both
have
a
value
of
8.8.
This
value
may
be
derived
from
another
phthalate,
which
makes
it
unreliable
toward
the
specific
phthalate
(ECPI
2014
and
Megaloid1
2013).
All
in
all
more
attention
should
be
placed
upon
sediment
as
it
tends
to
have
the
highest
levels,
even
during
different
seasons
(Figure
5)
(Sibali
et
al.,
2013).
All
phthalates
however
have
a
low
solubility
meaning
that
once
saturated
in
the
water,
particles
of
phthalate
will
join
the
sediment
(Sibali
et
al.,
2013).
Figure
5
shows
Figure
5:
Sediment
and
water
levels
of
phthalates
(DEHP,
DBP,
DEP
and
DMP
at
different
sample
sites
along
the
River
Jeksei
during
two
seasons
(Sibali
et
al.,
2013).

8. 8
the
differences
in
water
and
sediment
phthalate
levels
from
the
River
Jukskei,
South
Africa.
2.2.1
Differences
in
seasons
It
is
still
unclear
why
these
differences
in
seasons
arise.
For
atmospheric
phthalates
for
example
seasonal
differences
can
be
due
to
influences
of
emission
sources
such
as
the
burning
of
coal
in
cold
season
that
would
then
produce
phthalate
particulates
in
the
air
(Kong
et
al.,
2013).
Another
reason
could
be
a
meteorological
parameter.
Intense
sunlight
during
the
summer,
when
photochemical
reactions
are
increased
and
degrade
phthalates
lowering
the
concentrations
within
the
atmosphere.
Rain
can
also
be
a
culprit
through
diluting
and
washing
away
phthalates
particulates
(Kong
et
al.,
2013).
When
comparing
the
water
and
sediment
levels
in
the
graph
above
it
is
possible
that
the
high
winter
levels
are
due
to
a
lack
of
rain
therefore
concentrating
the
phthalates.
African
summer
(rain
period)
could
perhaps
dilute
the
phthalate
concentrations
within
the
water
and
sediment
therefore
lowering
the
concentrations
(Sibali
et
al.,
2013).
Plants
have
also
very
recently
been
shown
to
significantly
enhance
the
dissipation
of
phthalates
in
soil
in
three
ways:
phytoremediation,
increased
sorption
of
phthalates
to
soil
and
plant
promoted
biodegradation
(Li
et
al.,
2004).
This
could
be
another
explanation
for
the
lower
summer
concentrations
of
phthalates
in
figure
5.
Half-­‐lives
of
phthalates
can
also
be
increased
through
increased
sorption
and
cooler
temperatures
(Staples
et
al.,
1997
and
Kickham
et
al.,
2012).
2.3
Levels
in
the
environment
In
the
1990’s
the
levels
of
phthalates
in
river
water,
in
Manchester,
UK
for
example,
were
at
a
mean
of
21.5μg/L
±12.5
and
1.3μg/L
±0.9
for
DBP
and
DEHP
respectively
(Fatoki
and
Vernon,
1990).
High
standard
deviation
was
due
to
the
different
sample
station
along
the
river
Irwell.
However
surprisingly
levels
at
the
effluent
of
a
sewage
treatment
plant
were
the
lowest
at
6μg/L
for
DBP
while
all
other

9. 9
sample
sites
were
above
12.1μg/L.
For
DEHP
the
highest
concentration
was
found
at
the
sewage
treatment
plant
(1.9μg/L)
that
also
coincided
with
the
percentage
of
DEHP
found
in
the
samples
1.9%
for
DEHP
(79.4%
for
DBP).
This
contradicts
previous
research
claiming
that
DEHP
has
the
highest
environmental
levels.
However
this
could
be
due
to
the
higher
degradability
of
DEHP
under
anaerobic
conditions
(Huang
et
al.,
2008).
In
Germany
DEHP
surface
water
levels
ranged
between
0.33-­‐97.8μg/L
and
sediment
levels
varied
between
0.21-­‐8.44mg/kg
dry
weight
and
for
DBP
0.12-­‐
8.80μg/L
and
0.06-­‐2.08mg/kg
dry
weight,
respectively
(Fromme
et
al.,
2002).
This
study
showed
both
phthalates
to
have
a
wide
variability
in
levels
throughout
Germany
although
DEHP
always
had
the
highest
levels.
In
the
Netherlands
environmental
measurements
were
taken
in
2005
and
it
was
found
that
DEHP
showed
the
highest
concentrations
in
both
mean
municipal
sewage
treatment
plant
and
industrial
waste
water
types
(Vethaak
et
al.,
2005).
Mean
municipal
sewage
treatment
plant
effluent
levels
were
around
1.5μg/L
and
industrial
wastewater
levels
were
150μg/L
compared
to
DBP
that
showed
levels
of
0.3
and
2.2μg/L.
70%
of
the
DEHP
and
DBP
samples
contained
levels
above
the
level
of
detection
(LOD)
although
only
30%
of
the
DBP
samples
were
above
the
LOD
in
the
sewage
treatment
plant
effluent
water.
Fish
muscle
concentrations
of
Bream
(Abramis
brama)
and
Flounder
(Platichthys
flesus)
showed
mean
concentrations
of
0.044μg
DBP/g,
0.153μg
DEHP/g
and
0.0078μg
DBP/g,
0.064μg
DEHP/g
in
each
fish
respectively
(Vethaak
et
al.,
2005).
It
seems
that
phthalate
concentration
varies
not
only
within
country
or
city
but
also
within
micro
environments
and
water
types.
2.4
Half-­‐lives
Half-­‐lives
of
phthalates
are
the
time
for
a
substance
to
fall
to
half
its
original
concentration
(i.e.
degrading)
(Staples
et
al.,
21997)
through
hydrolysis
of
ester
bindings
(Liang
et
al.,
2008)
and
the
range
of
half-­‐
lives
referring
to
phthalates
is
vast.
Staples
et
al.,
(1997)
reported
a
half-­‐life
of
28
day
on
average
for
phthalates
within
sewage
sludge,

10. 10
while
within
the
atmosphere
half-­‐lives
consist
of
around
one
day
(DBP-­‐<6
days,
DEHP-­‐<2
days,
DINP-­‐<2
days).
Within
sediment
half-­‐
lives
of
approximately
<one
week
–
several
months
may
be
recorded
and
within
surface
waters
<one
day
–
two
weeks
(Staples
et
al.,
1997).
Staples
et
al.
also
reported
a
half-­‐life
of
years
through
aqueous
hydrolysis
(DBP–22
years,
DEHP-­‐2000
years).
In
contrast
Yuan
et
al.
(2010)
reported
that
DBP
and
DEHP
had
half-­‐lives
of
1.6-­‐2.9
days
and
5.0-­‐8.3
days
within
sediment,
respectively.
It
has
also
been
postulated
that
DEHP
degrades
fairly
rapidly
under
aerobic
conditions
(Brooke
et
al.,
1991).
Microbial
degradation
has
shown
DBP
to
be
completely
degraded
within
28
days
(Liang
et
al.,
2008).
In
Turner
and
Rawling
(2000)
eight
phthalates
were
found
in
a
water
sample
and
half-­‐lives
were
measured.
On
average
the
phthalate
half-­‐
life
in
aerobic
conditions
was
between
2.4-­‐14.8
days
and
14-­‐34
days
under
anaerobic
conditions.
Other
studies
such
as
Yuwatini
et
al.,
(2006)
showed
that
DEHP
half
life
in
water
is
approximately
two
days
while
in
sediment
it
can
last
up
to
14
days.
Magdouli
et
al.,
(2013)
stated
that
half-­‐lives
of
DEHP
are
<one
month
in
aerobic
conditions
and
>one
month
in
anaerobic
conditions.
In
water
(with
sun)
under
acidic
conditions
half-­‐lives
can
be
around
390
days
while
in
neutral
conditions
may
be
up
to
1600
days.
From
above
it
is
clear
that
phthalate
half-­‐lives
may
have
wide
ranges.
This
is
due
to
the
different
environmental
compartments
in
which
the
phthalate
may
be
present
i.e.
atmosphere,
sediment,
water,
inside
the
organism
as
each
situation
will
affect
the
half-­‐life
as
well
as
what
process
of
degradation
is
measured.
This
makes
it
difficult
to
consent
on
fixed
half-­‐life
values.
In
general
it
is
thought
that
the
longer
the
phthalate
chain
(R
group
in
figure
1)
the
longer
the
half-­‐life
and
the
more
persistent
it
will
be
and
that
aerobic
conditions
will
almost
most
certainly
speed
up
degradation
compared
to
anaerobic
(Liang
et
al.,
2008).
The
organization
for
economic
co-­‐operation
and
development
(OECD)
has
guideline
tests
and
criteria
for
defining
‘ready
biodegradability’.
Using
these
criteria,
>60%
removal
of
inorganic
carbon
within
a
10-­‐day
window
of
the
28-­‐day
test,
both
DBP
and
DEHP
are
readily
biodegradable
in
all
three
states
(water,
sediment
and
air).
Data
concerning
DINP
was
only
available
for

11. 11
atmospheric
half-­‐life
but
still
fits
within
the
criteria
for
bing
readily
biodegradable.
If
all
half-­‐life
tests
incorporated
these
test
guidelines
then
more
accurate
comparisons
could
be
made.
2.5
Inside
the
organism
Phthalate
accumulation
within
organisms
is
also
low,
partly
due
to
their
biodegradability
but
also
due
to
the
compound
itself
not
being
highly
accumulative
in
tissue,
rendering
phthalates
non
bio-­‐
accumulative
compounds
(Van
Den
Berg
et
al.,
2003;
Oehlmann
et
al.,
2009
and
Mankidy
et
al.,
2013).
Due
to
their
high
transformation
rate
phthalates
are
not
bio-­‐accumulative
(Mankidy
et
al.,
2013
and
Van
Den
Berg
et
al.
2003)
meaning
that
on
one
hand
the
phthalate
compound
is
transformed
into
a
metabolite
that
can
then
interact
with
receptors
and
enzymes
within
the
organism
(Euling
et
al.,
2013).
On
the
other
hand,
this
metabolism
also
produces
sulphates
and
other
glucuronides
that
assist
in
the
removal
of
the
parent
compound
(phthalate)
reducing
the
adverse
effects
of
the
phthalate
to
the
organism
and
also
through
the
food
chain
(Van
Den
Berg
et
al.,
2003
and
Van
Wezel
et
al.
2000).
2.6
Modes
of
action
once
inside
an
organism
Phthalates
being
EDC’s
have
a
multiple
array
of
modes
of
action
(MOA)
making
it
important
to
understand
how
the
EDC
interacts
on
a
cellular
level
(Nelson
and
Habibi,
2013).
Endogenous
hormones
(specifically
estrogen
and
androgen)
are
most
commonly
the
concern
when
regarding
phthalates.
Estrogenic
receptors
(ERs)
and
androgenic
receptors
(ARs)
are
important
in
reproduction
(ER
and
AR),
sexual
differentiation
(AR)
and
even
adult
sexual
behavior
(AR)
(Harbott
et
al.,
2007
and
Thibaut
and
Porte,
2004).
Peroxisome
proliferator
activated
receptors
(PPARs)
act
as
regulators
for
lipid
and
carbohydrate
metabolism
as
well
as
cell
differentiation
(Maradonna
et
al.,
2013).
Another
MOA
is
through
oxidative
damage
(OxD)
that
can
cause
disturbances
to
the
cellular
metabolism
(Harbott
et
al.,
2007).
All
these
receptors
are
present
on
cell
walls.
EDC’s
show
similar
biological
effects
to
estrogens
and
androgens
and
interfere
(agonistically/antagonistically)
with
the
cell
receptors
(Van

12. 12
den
Berg
et
al.,
2003)
either
decreasing
or
increasing
gene
expression,
production
of
hormones,
enzymes
and
phase
II
metabolites
affecting
the
level
of
active
hormones
present
within
an
organisms
(Thibaut
and
Porte,
2004).
2.7
Environmental
risk
limits
The
European
commission
previously
considered
the
four
phthalates
(DEHP,
DBP,
DINP
and
DIDP)
priority
substances
meaning
that
environmental
risk
assessments
(ERA)
must
have
been
carried
out
on
these
substances
(Oehlmann
et
al.,
2008).
ERA’s
compare
environmental
concentrations
or
predicted
environmental
concentrations
(PEC)
with
the
predicted
no-­‐effect
concentrations
(PNEC).
When
the
PEC/PNEC
ratio
is
<1
there
is
no
risk,
where
as
if
the
ratio
is
≥1
there
is
a
potential
risk
meaning
strategies
must
be
put
in
place
to
reduce
the
concentrations.
For
the
EPA
to
recognize
acute
effects,
a
total
of
five
tests
must
be
completed
on
at
least
four
different
species
using
the
limit
of
solubility
concentration
(max.
3μg/L)
(Oelmann
et
al.,
2008).
By
2004
the
European
union
risk
assessment
reports
stated
that
for
DBP,
DINP
and
DIDP
there
was
no
need
for
testing
or
information.
DEHP
was
not
granted
similar
status
and
therefore
still
remained
on
the
priority
substance
list
in
2008
(EC,
2014
and
Oehlmann
et
al.,
2008).
2.8
Objective
This
paper
will
focus
on
plastic
derived
EDC
known
as
phthalates.
Background
on
phthalates
and
why
they
are
the
focus
of
research
will
be
given.
It
will
highlight
the
associated
endocrine
disruptions
(developmental
and
reproductive)
and
will
speculate
to
future
work.
In
previous
reviews,
fish
have
never
been
the
sole
focus
neither
experiment
set
up
explained.
It
has
been
approximately
13
years
since
the
last
review
that
incorporated
over
12
studies
(Van
Wezel
et
al.,
2000).
The
paper
will
focus
on
four
phthalates
allowing
a
more
refined
and
in
depth
review.

13. 13
III.
Method
For
this
paper
focus
was
on
the
compounds
DEHP,
DBP
DINP
and
DIDP
due
to
their
high
abundance
within
the
environment
(former
two)
and
acclaimed
‘no
effects’
of
the
latter
two
(Oehlman
et
al.,
2009
and
Hallmark,
2010).
Searches
were
be
carried
out
on
’google’
‘google
scholar’
and
‘Web
of
science’
focusing
mainly
on
publications
within
the
years
2000-­‐2014.
Searches
for
DEHP,
DBP,
DINP,
DIDP,
effects
of
DEHP/DBP/DINP/DIDP
on
aquatic
organisms/fish,
reproductive/developmental/metabolic
effects
of
phthalates,
vitellogenin
effects
of
phthalates,
intersex
caused
by
endocrine
disrupters,
and
environmental
phthalates
are
a
few
of
the
search
terms
used.
The
main
duration
of
research
lasted
approximately
3
weeks-­‐1
month
and
only
full
text
articles
were
incorporated
within
the
paper.
IV.
Results
A
total
of
46
papers
were
gathered
and
divided
into
sections
on
organism
toxicity
(≈18),
phthalate
levels
in
the
environment
(≈5),
general
information,
however
nearly
all
articles
had
multiple
section
uses.
Due
to
the
majority
of
organism
toxicity
publication
a
further
division
of
pre
2000
and
post
2000
research
as
well
as
compound
groups
were
added.
This
was
done
due
to
the
majority
of
papers
found
being
post
2000
and
to
separate
‘recent
work’
from
‘previous
work’.
All
publications
were
given
in
publication
date
order
(oldest-­‐
newest).
Most
experiments
within
the
ecotoxicology
field
focus
on
either
in
vivo
or
in
vitro
set-­‐ups.
The
former
refers
to
the
whole
organism
being
studied
allowing
observation
of
the
overall
effect
of
compounds
on
the
organism.
The
latter
refers
to
using
cells
in
controlled
environments
(such
as
petri
dishes,
assays,
etc)
where
for
example
assays
can
provide
information
on
the
mechanism
of
action
(MOA)
of
certain
compounds;
unfortunately
this
does
not
mimic
the
whole
organism
(Sohoni
and
Sumpter,
1998).

14. 14
4.1
Summary
of
literature
(1980-­‐1999)
Table
1:
ED
effects
of
phthalates
in
in
vitro
receptor
binding
affinity
tests.
Cell
type
Effect
(mM)
Remark
Original
references
(within
Van
Wezel
et
al.,
2000)
DBP
Trout
hepatocyte
EC50
=
1
REP:
6.7x10-­‐6
Jobling
et
al.,
1995
Trout
hepatic
cytosol
EC
10-­‐25
=
0.17
REP:
2x10-­‐5
Knudsen
and
Pottinger,
1999
DEHP
Trout
hepatocyte
EC75
=
1
REP:
1x10-­‐5
Jobling
et
al.,
1995
Trout
hepatic
cytosol
EC10-­‐25
=
0.17
REP:
2x10-­‐5
Knudsen
and
Pottinger,
1999
DINP
Trout
hepatic
cytosol
No
effect
at
0.17
-­‐
Knudsen
and
Pottinger,
1999
REP:
relative
potency
compared
with
17-­‐estradiol
(Based
of
appendices
by
Van
Wezel
et
al.,
2000).
Table
2:
Toxicity
data
for
DBP.
1-­‐Y:
chemical
analyzed
in
test
solution
and
N:
chemical
not
analyzed
in
test
solution
or
no
data.
2-­‐S:
static,
R:
Static
with
renewal
and
F:
flow
through.
3-­‐S:
survival,
R:
reproductive
and
G:
Growth.
*-­‐Average
of
results
(mg/L)
when
all
parameters
and
authors
were
the
same
(Based
of
appendices
by
Van
Wezel
et
al.,
2000).
Organism
Chemical
analysis1
Test
type2
Exp.
time
End
point3
Results
(mg/L)
Original
references
(within
Van
Wezel
et
al.,
2000)
Chronic
toxicity
to
freshwater
organisms:
NOEC
values
Oncorhynchus
mykiss
Y
P
60d
G
0.1
Rhodes
et
al.,
1995
Pimephales
promelas
Y
F
20d
G
0.56
McCarthy
and
Whitmore,
1985
Acute
toxicity
to
freshwater
organisms:
L(E)C50
values
Brachydanio
rerio
Y
S,
R
96h
S
2.2
Scholz,
1994
Lepomis
macrochirus*
N
S
96h
S
1.6
Mayer
and
Ellersieck,
1986
Lepomis
macrochirus
N
F
96h
S
1.6
Mayer
and
Ellersieck,
1986
Oncorhynchus
mykiss*
N
S
96h
S
4.4
Mayer
and
Ellersieck,
1986
Oncorhynchus
mykiss
N
F
96h
S
1.5
Mayer
and
Ellersieck,
1986
Oncorhynchus
mykiss
Y
-­‐
96h
S
1.2-­‐1.8
Hrudey
et
al.,
1976
Oncorhynchus
mykiss
Y
F
96h
S
1.6
Adams
et
al.,
1995
Perca
flavescens
N
F
96h
S
0.35
Mayer
and
Ellersieck,
1986
Pimephales
promelas
N
S
96h
S
1.3
Mayer
and
Ellersieck,
1986
Pimephales
promelas
N
F
96h
S
4
Mayer
and
Ellersieck,
1986
Pimephales
promelas
N
-­‐
96h
S
2
McCarthy
and
Whitmore,
1985
Pimephales
promelas
Y
S
96h
S
1.5
Adams
et
al.,
1995
Pimephales
promelas*
Y
F
96h
S
0.97
DeFoe
et
al.,1990
Pimephales
promelas
Y
F
96h
S
0.92
Adams
et
al.,
1995

15. 15
Table
3:
Toxicity
data
for
DEHP.
Organism
Chemical
analysis1
Test
type2
Exp.
time
End
point3
Results
(mg/L)
Original
references
(within
Van
Wezel
et
al.,
2000)
Chronic
toxicity
to
freshwater
organisms:
NOEC
values
Brachydanio
rerio
N
R
35d
S,
G
≥0.32
Canton
et
al.,
1984
Gasterosteus
aculeatus
N
-­‐
28d
S,
G
≥0.32
Van
den
Dikkenberg
et
al.,
1989
Jordanella
floridae
N
S
28d
S,
G
≥0.32
Adema
et
al.,
1981
Oncorhynchus
mykiss
Y
F
102d
S,
R
0.005
Mehrle
and
Mayer,
1976
Oncorhynchus
mykiss
Y
F
90d
S,
G,
R
>0.5
DeFoe
et
al.,
1990
Oncorhynchus
mykiss
Y
F
70d
S,
G,
R
>0.0073
Cohle
and
Stratton,
1992
(EU
draft)
Oryzias
latipes
Y
F
168d
G
0.55
DeFoe
et
al.,
1990
Oryzias
latipes
N
R
28d
S,
G
≥0.32
Adema
et
al.,
1981
Pimephales
promelas
Y
F
56d
S,
G
0.062
Mehrle
and
Mayer,
1976
Poecilia
reticulata
N
-­‐
28d
S,
G
≥0.32
Adema
et
al.,
1981
Acute
toxicity
to
freshwater
organisms:
L(E)C50
values
Brachydanio
rerio
N
-­‐
96h
S
>0.32
Van
den
Dikkenberg
et
al.,
1989
Brachydanio
rerio
Y
R
96h
S
>100
Scholz,
1995
Gasterosteus
aculeatus
N
-­‐
96h
S
>0.32
Van
den
Dikkenberg
et
al.,
1989
Ictalurus
punctatus
-­‐
S
96h
S
>10
Mayer
and
Sanders,
1973
Ictalurus
punctatus
Y
F
96h
S
>100
Johnson
and
Finley,
1980
Ictalurus
punctatus
N
S
24h
S
>100
Mayer
and
Ellersieck,
1986
Ictalurus
punctatus
N
S
96h
S
>100
Mayer
and
Ellersieck,
1986
Ictalurus
punctatus
N
F
96h
S
>0.2
Mayer
and
Ellersieck,
1986
Jordanella
floridae
N
-­‐
96h
S
>0.32
Van
den
Dikkenberg
et
al.,
1989
Lepomis
macrochirus
-­‐
S
96h
S
>10
Mayer
and
Sanders,
1973
Lepomis
macrochirus
N
S
96h
S
>250
Bionomics
Inc.,
1972
Lepomis
macrochirus
Y
F
96h
S
>100
Johnson
and
Finley,
1980
Lepomis
macrochirus
Y
S
96h
S
>0.2
Adams
et
al.,
1995
Lepomis
macrochirus
N
S
24h
S
>100
Mayer
and
Ellersieck,
1986
Lepomis
macrochirus
N
S
96h
S
>100
Mayer
and
Ellersieck,
1986
Lepomis
macrochirus
N
F
96h
S
>0.2
Mayer
and
Ellersieck,
1986
Oncorhynchus
mykiss
-­‐
S
96h
S
>10
Mayer
and
Sanders,
1973
Oncorhynchus
mykiss
-­‐
S
96h
S
>1000
Silvo,
1974
(EU
draft)
Oncorhynchus
mykiss
N
S
96h
S
>540
Hrudey
et
al.,
1976
Oncorhynchus
mykiss
Y
F
96h
S
>0.32
Adams
et
al.,
1995
Oncorhynchus
kisutch
N
S
24h
S
>100
Mayer
and
Ellersieck,
1986
Oncorhynchus
kisutch
N
S
96h
S
>100
Mayer
and
Ellersieck,
1986
Oncorhynchus
mykiss
N
S
24h
S
>100
Mayer
and
Ellersieck,
1986
Oncorhynchus
mykiss
N
S
96h
S
>100
Mayer
and
Ellersieck,
1986
Oncorhynchus
mykiss
Y
F
96h
S
>20
DeFoe
et
al.,
1990
Oryzias
latipes
N
-­‐
96h
S
>0.32
Van
den
Dikkenberg
et
al.,
1989
Oryzias
latipes
Y
F
96h
S
>0.67
DeFoe
et
al.,
1990
Pimephales
promelas
-­‐
S
96h
S
>10
Mayer
and
Sanders,
1973
Pimephales
promelas
Y
F
96h
S
>0.67
DeFoe
et
al.,
1990

16. 16
Pimephales
promelas
N
F
96h
S
>1
Mayer
and
Ellersieck,
1986
Pimephales
promelas
Y
F
96h
S
>0.33
DeFoe
et
al.,
1990
Pimephales
promelas
Y
S
96h
S
>0.16
Adams
et
al.,
1995
1-­‐Y:
chemical
analyzed
in
test
solution
and
N:
chemical
not
analyzed
in
test
solution
or
no
data.
2-­‐S:
static,
R:
Static
with
renewal
and
F:
flow
through.
3-­‐S:
survival,
R:
reproductive
and
G:
Growth.
EU
draft:
(DEHP)
(Based
of
appendices
by
Van
Wezel
et
al.,
2000).
4.2
Literature
(1980-­‐1999)
A
meta-­‐analysis
using
journals
from
1980-­‐1999
was
carried
out
by
Van
Wezel
et
al.,
2000.
Above
in
table
1,
2
and
3
a
summary
of
these
results
(no
observed
effect
concentration-­‐NOEC,
X%
effective
concentrations-­‐ECx,
chronic
and
acute
exposure)
concerning
fish
can
be
found.
They
found
minimal
difference
between
nominal
and
actual
concentrations
used
in
studies
concerning
DBP.
The
most
sensitive
freshwater
organism
was
Oncorhynchus
mykiss
that
showed
the
lowest
chronic
NOEC
at
0.1mg/L
(table
2).
Acute
toxicity
data
was
more
available
(see
table
5)
and
Van
Wezel
et
al.
reported
that
‘no
useful
test’
regarding
soil
or
sediment
was
found.
When
comparing
chronic
and
acute
DEHP
results
it
was
found
that
both
categories
showed
no
effects
in
the
majority
of
the
studies
(even
at
the
highest
concentration
tested
acute:
0.55mg/L
and
chronic:
1x106mg/L).
When
effects
were
recorded
and
NOEC
could
be
produced
the
NOEC
was
above
the
water
solubility
of
phthalates
(3μg/L).
With
all
the
data
available
the
authors
derived
an
ERL
for
the
aquatic
and
sediment
environments.
For
DBP
this
was
done
by
using
the
lowest
NOEC
(0.1mg/L)
and
applying
an
assessment
factor
of
10.
For
sediment
due
to
lack
of
data
the
ERL
was
derived
by
multiplying
the
lowest
Koc,
partition
coefficient
between
organic
carbon
in
the
soil/sediment
and
water,
value
(1.2x103L/kg:
12mg/kg).
For
DEHP,
due
to
no
effects
observed,
the
NOEC
for
the
only
soil
organisms
(Rana
arvalis
–
frog)
was
used
10mg/kg
fresh
weight
and
applying
a
factor
of
10.
The
ERL
for
soil
was
then
used
to
derive
an
ERL
for
water
by
combining
with
the
lowest
soil/sediment
Koc.
The
derived
ERLs
for
DBP:
10μg/L
and
0.7mg/Kg
fresh
weight
and
DEHP:
0.19μg/L
and
1.0mg/Kg
fresh
weight.
When
surface
water

17. 17
samples
were
taken
at
different
location
in
the
Netherlands
they
found
that
DBP
levels
were
rarely
above
the
ERL
(both
water
and
sediment)
derived
in
this
study.
For
DEHP
however
unexpected
levels
3-­‐20
times
higher
than
the
derived
ERL
for
water
were
observed
and
sediment
levels
were
also
much
higher
than
the
derived
sediment
ERL.
4.3
Summary
of
literature
post
2000
Table
4:
DBP
summary.
N-­‐Depicts
nominal
concentrations.
A-­‐depicts
acute
exposure
studies.
C-­‐depicts
chronic
exposure
studies.
[]-­‐concentration
causing
significant
effects.
VTG-­‐vitellogenin.
D/hpf-­‐days/hours
post
fertilization.
Dep.-­‐depuration
(none
contaminated
water).
*-­‐<0.05,
**<0.01,
***<0.001
significance
levels.
Species
Age,
sex,
exp.
type
and
concentration
(μg/L)
(unless
indicated)
Exposure
route
and
duration
Effects
Authors
Sander
lucioperca
Juvenile
(61
dph)
(in
vivo)
0.125,
0.25,
0.5,
1,
2g/Kg
feed
Food
5
weeks
*No
effects
on
female
fish,
growth
rate
and
survival.
*Increases
in
[DBP]
shows
decreases
in
male
specimens.
Jarmolowic
z
et
al.,
(2003)
Danio
rerio
Adult
male
(in
vivo
and
in
vitro)
500
Water
15
days
Day
7:
*increase
in
surface
density
of
peroxisomes.
Day
15:
*increase
in
both
surface
density
and
numerical
density
of
peroxisomes.
Increase
in
activity
of
acyl-­‐CoA
oxidase.
Ortiz-­‐
Zarragoitia
and
Cajaraville
(2005)
Danio
rerio
1)
Embryos
(1-­‐2
hpf)
25,
100
2)
Adult
female
100,
500
(in
vivo)
Water
1:
8
weeks
2:
15
days
1)
[100]
Increase
in
number
and
volume
of
peroxisome
density
and
acyl-­‐CoA
oxidase
enzyme.
2)
Mortality
of
female
offspring
increased.
Ortiz-­‐
Zarragoitia
et
al.,
(2006)
Gasterosteus
aculetaus
Adult
male
(in
vivo)
N50,
N100
(Measured
levels
15,
35
respectively)
Water
22
days
[35*/**]
Increase
in
testosterone
and
oxidised
testosterone.
Decrease
in
spiggin
(protein
glue).
Aoki
et
al.,
(2011)
Cyprinus
carpio
(in
vitro)
100μM,
1mM
Incubated
in
vitro
[100]
Inhibited
formation
of
5α-­‐
Adione
and
synthesis
of
5α-­‐DHT.
[1]
Increased
synthesis
of
17α,20α/βDP
Thibaut
and
Porte
(2004)
Pimephales
promelis
1
hpf
(in
vivo
and
In
vitro)
1000
Water
96
hpf
None
Mankidy
et
al.,
(2013)

18. 18
Table
5:
DEHP
summary.
N-­‐Depicts
nominal
concentrations.
[]-­‐concentration
causing
significant
effects.
Hpf-­‐
hours
post
fertilization.
Dph-­‐days
post
hatch.
Dep.-­‐depuration
(none
contaminated
water).
*-­‐<0.05,
**<0.01
significance
levels.
Species
Age,
sex,
exp.
type
and
concentration
(μg/L)
(unless
indicated)
Exposure
route
and
duration
Effects
Authors
Oryzias
laptipes
Adult
male
(in
vivo)
N0.1,
N0.3,
N1μmol
Water
2
weeks
None
Shioda
and
Wakabayashi
(2000)
Oryzias
laptipes
A)
a
few
days
old
10,
50,
100
C)
7
month
N1,
N10,
N50
(in
vivo
and
in
vitro)
Water
A:
5
days
C:
3
months
A)
[all]
VTG
protein
not
present
in
males.
C)
[N10*,
N50]*
GSI
lower
in
females
and
[Nall]
retardation
in
ovary
(oocyte)
development.
Kim
et
al.,
(2002)
Oryzias
laptipes
1dpf
(in
vivo)
N0.01,
N0.1,
N1,
N10
Water
Until
hatched
[N0.1,
N1]
Decreased
hatch
time.
Post
5-­‐
6
months
dep.:
[N0.01***,
N0.1*,
N1***]
increased
mortality,
[N0.01*]
altered
sex
ratio.
[N0.1*,
N1*,
N10**]
decrease
in
male
body
weight.
Chikae
et
al.,
(2004)
Poecilia
reticulate
<1
week
(in
vivo)
0.1,
1,
10
Water
3
months
Day
14:
[10]
decrease
in
length
and
weight.
Day
49
and
91:
[1,
10]
decrease
in
length
and
weight
(more
significant
in
females
[all]**)
Zantonelli
et
al.,
(2009)
Danio
rerio
6
month
female
(in
vivo
and
In
vitro)
0.02,
0.2,
2,
20,
40
Water
3
weeks
[2]
Increase
in
VTG
oocytes
and
decrease
in
pre-­‐VTG
oocytes*.
[all]
down
regulation
of
LHR
and
plasma
VTG.
Carnevali
et
al.,
(2010)
Danio
rerio
Mature
males
(in
vivo)
0.5,
50,
5000
mg/kg
body
weight
Injection
on
day
1
and
5
[5000***]
Decrease
in
fertilization
success.
[50*,
5000**/***]
decrease
in
no.
of
spermatozoa
and
increase
in
no.
of
spermatocytes.
[5000*]
increase
in
VTG
levels
(males
should
not
have
VTG).
[5000**/*]
increase
expression
of
acox1
and
ehhadh.
Uren-­‐
Webster
et
al.,
(2010)
Cyprinus
carpio
(in
vitro)
100μM,
1mM
Incubated
in
vitro
[100]
Inhibited
formation
of
5α-­‐Adione.
Thibaut
and
Porte
(2004)
Pimephales
promelis
1
hpf
(in
vivo
and
In
vitro)
1000
Water
96
hpf
[1000**/*]
increase
in
embryo
mortality
and
increased
lipid
peroxidation
Mankidy
et
al.,
(2013)
(Salmo
salar)
4
weeks
post
hatch
(in
vivo)
N0,
N400,
N800,
N1500mg/kg
feed
Food
4
weeks
[DEHP]
3x
greater
then
MEHP
(metabolite).
1
week
after
dep.:
DEHP+MEHP
levels
returned
to
background
levels.
1
month
dep.:
[N1500*]
ovo-­‐testis
combo
Norman
et
al.,
(2007)

19. 19
Table
6:
DINP
and
DIDP
summary.
Species
Age,
sex,
exp.
type
and
concentration
(μg/L)
(unless
indicated)
Exposure
route
and
duration
Effects
Authors
Oryzias
laptipes
2
week
old
larvae
(in
vivo)
20μg/g
(for
both
DINP
and
DIDP)
Food
F0
till
gen.
F2-­‐42dph
DINP:
F0
Embryo
development
showed
decreases
in
red
blood
cell
pigment*.
DINP:
F1
survival
decreased*
Patyna
et
al.,
(2006)
Dph-­‐days
post
hatch.
*-­‐<0.05
significance
levels.
Patyna
et
al.
2006
uses
two
bioassays
for
a
single
effect.
Therefore
only
effects
proving
significant
on
both
assays
are
used
in
this
table.
4.4
Literature
post
2000
4.4.1
DEHP
In
2000
Shioda
and
wakabayashi
studied
the
effects
of
DEHP
(in
vivo)
on
the
number
of
eggs
produced
by
mating
couples
and
number
of
successful
hatchings
in
medaka
fish
(Oryzias
latipes).
For
this
experiments
groups
(one
male
and
two
females)
with
the
highest
number
of
fertilized
eggs
were
used.
Males
were
exposed
for
two
weeks
to
low
nominal
DEHP
concentrations
of
0.1,
0.3
and
1μmol/L
(through
means
of
water)
along
side
a
positive
(17
β–estradiol,
a
natural
estrogen)
and
negative
control
(tap
water).
Once
exposed
the
males
were
placed
back
into
their
original
group.
The
DEHP
concentrations
showed
no
significant
effects
on
number
of
eggs
and
hatchings,
which
could
be
due
to
the
extremely
low
concentrations
used.
Both
chronic
and
acute
exposures
of
DEHP
(in
vivo
and
in
vitro)
were
studied
by
Kim
et
al.
in
2002.
Japanese
Medaka
fish
(seven
months
old)
were
exposed
via
water
to
concentration
of
10,
50
and
100μg/L
of
DEHP
(for
acute
testing).
For
chronic
testing
fish
a
couple
days
old
were
exposed
to
nominal
concentrations
of
1,
10
and
50μg/L.
In
acute
exposure
(5
days)
it
was
found
that
the
protein
(200-­‐kDa)
used
for
identification
of
vitellogenin
(VTG)
proteins
were
not
present
in
male
Medaka
in
all
four
exposures
(including
the
control).
In
females
however
VTG
was
found
in
all
the
control
and
the
exposed
fish,
although
two
out
of
the
five
fish
in
1μg/L
exposed
tank
showed

20. 20
extremely
low
levels.
Overall,
acute
effects
of
DEHP
on
VTG
were
not
significant.
The
chronic
exposure
(three
months)
to
DEHP
showed
the
200-­‐kDa
protein
not
to
be
present
in
male
fish.
In
females
fish
however
the
protein
occurred
less
frequently
as
DEHP
concentration
increased.
The
weight
and
length
of
fish
used
in
the
chronic
exposure
showed
no
statistical
difference
in
all
treatments
showing
DEHP
to
have
no
effect
on
growth.
The
Gonado-­‐somatic
index
(GSI)
of
females
in
both
10
and
50μg/L
DEHP
treatments
was
statistically
lower
than
that
of
the
control
females
while
no
effect
was
found
on
male
fish
showing
DEHP
to
inhibit
the
development
of
Medaka
fish
ovaries.
Histology
of
both
the
gonads
and
ovaries
from
the
chronically
exposed
fish
were
also
looked
at.
Here
gonads
of
the
male
fish
were
not
deformed
compared
to
the
control,
while
the
oocytes
within
the
ovaries
of
female
fish
were.
In
the
control
females,
oocytes
were
developed
to
either
stage
two
or
three
(stage
three
allowing
them
to
be
fertilized).
In
all
1,
10
and
50μg/L
DEHP
treatments
only
37%,
0%
and
22%,
respectively,
of
the
fish
had
matured
oocytes
at
stage
three
compared
to
54%
of
the
control
–
taking
note
that
10μg/L
showed
no
stage
three
development.
Along
side,
only
26%,
25%
and
12%
of
the
female
fish
(respectively
of
1,
10,
50μg/L
DEHP)
could
reach
stage
one
compared
to
the
control
where
oocytes
development
was
not
stopped
(figure
2).
This
shows
the
retardation
effects
in
ovary
growth
of
DEHP
using
environmentally
relevant
concentrations.
In
2004,
Chikae
et
al.
also
conducted
an
in
vivo
study
on
the
negative
(irreversible)
effects
that
DEHP
exposure
using
pre-­‐hatched
Medaka
would
have
on
adulthood
(5-­‐6
months
post
hatch).
Treatments
of
water
containing
nominal
DEHP
concentrations
of
0.01,
0.1,
1,
10
Figure
2:
Ovaries
of
females
medaka
after
3
months.
A)
control,
developed
to
stage
3
B)
DEHP
(10μg/L)
stuck
in
stage
1
(Kim
et
al.,
2002)

21. 21
μg/L
and
a
control
were
used
to
expose
1-­‐day-­‐old
fertilized
eggs.
Once
hatched
the
fish
were
transferred
to
DEHP
free
water
for
5-­‐6
months.
At
the
beginning
(pre-­‐hatch)
over
90%
of
the
eggs
in
each
treatment
showed
signs
of
eye
development
(eyeing)
except
at
10μg/L
were
only
83%
were
found
eyeing.
Of
those
eggs
that
had
successful
eyeing,
over
90%
continued
to
hatch
in
each
treatment.
The
only
significant
difference
was
a
decrease
in
hatching
time
seen
at
the
0.1μg/L
(P<0.005)
and
1μg/L
DEHP
treatments
compared
to
the
control.
In
adulthood,
after
no
DEHP
exposure
for
5-­‐6
months,
irreversible
effects
were
significant
compared
to
the
control.
Post-­‐
hatch
mortality
was
significantly
increased
in
the
0.01,
0.1
and
1μg/L
treatments
(P<0.001,
<0.05
and
<0.001,
respectively).
Sex
ratio
within
the
0.01μg/L
treatment
was
significantly
altered
(4m:16f),
which
may
have
been
due
to
increased
male
mortality
or
feminization.
Body
weight
was
significant
different
in
male
fish
within
the
treatment
0.1,
1μg/L
(P<0.05)
and
10μg/L
(P<0.01).
This
study
shows
the
irreversible
effects
of
phthalate
exposure
in
embryonic
states
of
medaka
fish.
Norman
et
al.,
(2007)
studied
DEHP
(in
vivo)
on
Atlantic
salmon
(Salmo
salar)
with
nominal
concentrations
of
0,
400,
800
and
1500mg
DEHP/kg
feed.
Here
levels
of
DEHP
and
its
metabolite
mono-­‐
ethylhexyl
phthalate
(MEHP)
within
fish
tissue
were
studied
after
acute
exposure
(four
weeks)
of
DEHP.
Along
side,
histological,
growth
and
liver
effects
were
analyzed
after
one
month
of
depuration
(no
exposure
to
DEHP).
The
DEHP
concentration
in
the
fish
tissue
post
acute
phase
was
three
times
higher
than
the
concentration
of
MEHP.
Control
fish
that
were
not
Figure
3:
guppy
fish
at
day
49
with
treatments
above.
Grid
is
1mm
(Zanotelli
et
al.,
2009).

22. 22
exposed
to
dietary
DEHP
showed
low
background
levels
of
DEHP
(0.016
mg/kg
fish)
and
MEHP
(0.020
mg/kg
fish).
DEHP
and
MEHP
concentrations
increased
in
tissue
as
treatment
concentration
increased.
Both
were
eliminated
to
near
background
levels
one
week
after
the
depuration
phase.
Mortality
in
all
groups
was
low
(4%)
and
no
difference
in
weight
and
sex
ratio
was
recorded
between
the
different
exposure
concentrations.
Within
each
treatment
a
few
fish
(1%
of
400
and
1500mg
DEHP/kg
food)
were
observed
anatomically
to
be
slightly
different
(increased
testes
size).
The
only
statistically
difference
recorded
was
in
the
treatment
group
of
1500mg
DEHP/kg
feed
where
6
out
of
the
202
fish
had
ovo-­‐testis
(P<0.014).
This
study
showed
that
DEHP
had
no
short-­‐term
effects.
Zanotelli
et
al.,
(2009)
conducted
a
study
focusing
on
the
growth
(weight
and
length)
of
<1-­‐week-­‐old
(larval)
guppy
fish
(Poecilia
reticulata).
The
guppy
fish
were
subjected
to
continuous
exposure
(in
vivo)
to
DEHP
through
water
(0.1,
1,
10μg/L).
By
day
14
a
statistically
significant
growth
inhibition
at
the
highest
DEHP
concentration
was
observed
and
increased
with
time.
After
49
days
of
exposure,
DEHP
treated
fish
were
compared
to
control
fish.
Length
showed
a
dose-­‐
dependent
decrease,
where
DEHP
exposed
fish
at
1
and
10μg/L
were
15-­‐30%
shorter
(respectively)
than
the
control
and
weight
was
decreased
by
as
much
as
40-­‐70%
respectively.
After
91
days
of
chronic
exposure
to
environmentally
relevant
DEHP
concentrations
the
fish
showed
a
significant
decrease
in
weight
and
length:
fish
exposed
to
1
and
10μg/L
decreased
10%
and
26%
in
length
and
32
and
61%
in
weight,
respectively
(figure
3
below).
There
was
a
higher
level
of
significance
within
females
at
day
49,
with
all
concentrations
showing
a
P<0.01,
where
as
with
male
fish
at
day
49
only
10μg
DEHP/L
differed
form
the
control
with
P<0.01.
This
study
shows
that
chronic
exposure
as
low
as
1μg
DEHP/L
show
a
time
and
dose
dependent
relationship
when
it
comes
to
growth.
The
fish
used
in
this
study
were
considerably
small
which
could
have
increased
the
effects
observed.
Carnevali
et
al.,
(2010)
experimented
on
the
effects
of
DEHP
using
six-­‐month-­‐old
female
zebrafish
(Danio
rerio)
in
an
in
vivo
and
in
vitro

23. 23
study.
Environmentally
relevant
concentrations
of
0.02,
0.2,
2,
20,
40μg/L
as
well
as
a
positive
control
were
used
to
study
the
impact
on
fecundity,
ovulation
and
oocytes
maturation.
Fish
were
exposed
through
water
to
DEHP
for
three
weeks
and
were
compared
to
a
solvent
control.
Results
showed
that
fish
exposed
to
2μg/L
had
a
significant
increase
in
the
number
of
vitellogenic
oocytes.
This
was
associated
with
the
significant
decrease
in
pre-­‐vitellogenic
oocytes
compared
to
the
control
(P<0.05).
Down
regulation
of
ovarian
luteinizing
hormone
receptor
(LHR)
and
plasma
VTG
were
significantly
different
compared
to
the
control
at
all
five
doses
of
DEHP.
These
two
factors
clearly
show
the
estrogenic
activity
of
DEHP
with
regards
to
the
inhibition
of
oocytes
maturation.
This
is
also
supported
by
the
dose
dependent
increase
of
BMP15,
a
protein
involved
in
oocytes
maturation.
After
the
three-­‐week
exposure
period
the
female
fish
were
placed
into
a
mating
tank
with
control
males,
showing
that
the
fecundity
of
embryos
was
severely
compromised
compared
to
the
control.
This
study
shows
the
concrete
risk
associated
with
aquatic
organisms
living
in
phthalate-­‐
polluted
areas.
Another
in
vivo
and
in
vitro
experiment
on
DEHP
by
Uren-­‐Webster
et
al.,
(2010)
studied
the
reproductive
health
of
male
zebra
fish.
16
colonies
(male
and
female
pairs)
were
used
that
were
consistent
with
egg
production
and
spawning
were
over
a
10
day
period.
Here
instead
of
the
dietary
or
water
exposure
as
previous
studies
applied,
the
DEHP
solution
was
injected
into
the
intraperitoneal
cavity.
This
method
of
administration
allowed
all
fish
to
receive
the
same
dose
as
well
as
being
able
to
target
male
specimens.
Environmentally
relevant
concentrations
of
0.5mg
DEHP/kg
of
body
weight
(bw),
range
within
measured
concentration
of
wild
fish,
50mg/kg
bw
and
an
extremely
high
5000mg/kg
bw
was
used
to
assess
the
mechanisms
of
phthalate
toxicity.
All
three
treatments
were
compared
to
a
control.
The
fertilization
success
of
males
subjected
to
5000mg/kg
bw
were
significantly
lower
than
the
other
three
treatments
(P<0.001),
although
this
was
only
when
including
the
full
10
day
exposure
period
(the
first
5
day
period
showed
no
significant
difference).
No
abnormal
embryo
development
or
embryo
survival

24. 24
effects
were
seen
in
the
treatments.
Histological
analysis
of
the
gonads
showed
significantly
lower
numbers
of
spermatozoa
(sperm
cell)
in
the
testes
of
males
injected
with
50mg/kg
of
bw
(P<0.05)
and
5000mg/kg
bw
(P<0.01)
compared
to
the
control
fish.
On
the
other
hand
there
was
a
significant
increase
in
the
number
of
spermatocytes
(immature
male
germ
cell)
compared
to
the
control
in
both
50mg/kg
bw
(P<0.05)
and
5000mg/kg
bw
(P<0.001).
When
studying
at
the
liver,
a
statistically
significant
increase
(P<0.05)
in
VTG
levels
was
recorded
in
the
treatment
5000mg/kg,
which
showed
DEHP
to
have
estrogenic
activity,
as
VTG
should
not
be
found
in
male
zebra
fish.
In
the
male
fish
a
significant
increase
in
the
expression
of
the
genes
acox1
(acyl-­‐coenzyme
A
oxidase
1)
and
ehhadh
(enoyl-­‐coenzyme
A
hydratase/3-­‐hydroxyacyl
coenzyme
A
dehydrogenase)
that
are
both
involved
in
lipid
metabolism
was
found.
Males
showed
no
alterations
in
swimming
and
feeding
behavior
throughout
the
study
(compared
to
controls).
This
study
used
mature
fish
which
are
known
to
be
less
sensitive
than
juvenile
fish,
which
may
have
caused
the
conclusion
that
DEHP
at
environmentally
relevant
concentrations
(0.5mg
DEHP/kg
bd)
show
no
short
term
reproductive
effect.
Lee
and
Liang
(2011)
studied
zebra
fish
offspring
and
exposed
them
for
3
months
to
low
doses
of
DEHP
through
water
in
vivo.
2ml
of
DEHP
was
placed
into
tanks
containing
110
liters
of
water,
and
every
month
an
additional
0.1ml
of
DEHP
was
added.
They
observed
that
DEHP
altered
the
sex
ratio
from
1:1
to
3:7,
although
they
failed
to
specify
if
this
was
significant.
Decreases
in
growth
(length
and
weight)
were
observed,
but
were
however
not
significant.
They
concluded
that
DEHP
showed
no
effect.
4.4.2
DBP
In
Jarmolowicz
et
al.,
(2003)
DBP
concentrations
of
0.125,
0.25,
0.5,
1
and
2g/Kg
feed
were
used
to
determine
the
impact
on
the
reproductive
system
in
juvenile
European
pikeperch
(Sander
lucioperca)
in
an
in
vivo
study.
A
total
of
40
fish
were
placed
into
each
concentration
tank
with
a
control
tank
with
no
addition
of
DBP.
The

25. 25
experiment
was
divided
over
two
five
week
periods
the
first
being
61-­‐96
days
post
hatch
and
the
second
97-­‐132
days
post
hatch.
In
the
first
period
fish
were
fed
the
DBP
contaminated
feed.
During
the
second
period
fish
were
fed
uncontaminated
feed.
15
fish
from
each
tank
were
taken
for
histological
analysis
at
the
beginning
(60
days
post
hatch),
after
the
1st
and
the
2nd
period.
There
were
no
negative
changes
within
female
fish,
nor
in
survival
and
growth
rates
(P<0.05).
After
96
days
post-­‐hatch
the
sex
ratio
in
treatment
groups
0.125
and
025g/Kg
feed
was
1:1.
50%
of
the
males
in
those
two
groups
showed
gonads
that
were
comparable
to
those
of
the
control
group.
The
remaining
50%
showed
smaller
testes
size,
reduced
spermatogonia
(any
cell
of
the
gonad
which
matured
form
a
spermatocytes)
and
seminal
vesicles.
Increasing
concentration
of
DBP
showed
a
positive
correlation
with
reduction
in
male
specimens
(P<0.05).
Fish
within
the
treatment
group
2g/Kg
of
feed
had
a
significantly
altered
sex
ratio
(P<0.05).
In
the
two
highest
DBP
concentration
tanks
(1
and
2g/Kg
of
feed)
intersex
specimens
(6.7%)
were
recorded
although
not
significant.
Jarmolowicz
et
al.
concluded
that
DBP
acts
as
an
anti-­‐
androgen
(blocking
endogenous
androgen
action)
creating
an
‘estrogenic
environment’.
This
study
is
the
first
to
report
DBP
disruption
in
sex
differentiation
in
fish.
Ortiz-­‐Zarragoitia
and
Cajaraville
(2005)
used
high
DBP
concentrations
of
500μg/L
to
observe
effects
on
the
liver
peroxisomes,
enzyme
activity
of
Acyl-­‐CoA
oxidase
and
on
VTG
levels
(In
vivo
and
in
vitro).
They
exposed
adult
male
zebra
fish
through
water
for
15
days.
They
found
that
at
day
seven
the
surface
density
of
liver
peroxisomes
had
significantly
increased
(P<0.05)
compared
to
the
control
while
at
day
15
both
surface
density
and
numerical
density
had
significantly
increased
from
the
control
(P<0.05).
Acyl-­‐
CoA
oxidase
showed
a
significant
increase
in
activity
at
both
time
points
(days
7
and
15).
Surprisingly
DBP
showed
no
significant
effect
on
VTG
levels.
They
concluded
that
DBP
shows
no
estrogenic
effect
in
male
zebra
fish.

26. 26
The
next
year
(2006)
Ortiz-­‐Zarragoitia
et
al.,
conducted
another
study
(in
vivo)
on
DBP
and
the
Actyl-­‐CoA
oxidase
enzyme,
peroxisomes
and
VTG,
but
also
mortality.
This
study
was
conducted
in
two
parts,
the
first
focusing
on
early
life
exposure
and
the
second
focusing
on
adult
life
exposure
and
their
offspring.
For
the
first
experiment
zebra
fish
eggs
were
exposed
(via
water)
to
concentrations
of
25
and
100μg/L.
A
solvent
control
was
used
to
compare
results.
1-­‐2
hpf
eggs
were
exposed
for
three
weeks.
Once
hatched
they
were
transferred
to
a
larger
tank
and
exposed
for
a
further
five
weeks.
Measurements
were
taken
at
4,
6,
10
days
post
fertilization
(dpf)
and
3
and
5
weeks
post
fertilization
(wpf).
Results
showed
that
survival
of
exposed
fish
did
not
differ
from
the
controls.
However
anatomical
deformities
were
observed
in
both
DBP
exposed
groups
(figure
4).
Spinal
cord
malformations
and
hypertrophy
of
the
yolk
sack
were
noticed
in
infant
fish
and
in
juvenile
fish
spinal
cord
and
swim
bladder
malformations
were
apparent.
Although
Ortiz-­‐Zarragoitia
et
al.,
(2006)
fail
to
specify
numbers
of
malformed
fish,
however
those
in
the
control
showed
no
signs
of
malformation.
As
with
the
prior
study
in
2005,
here
too
they
found
that
the
number
and
volume
of
peroxisome
density
as
well
as
the
Acyl-­‐CoA
oxidase
enzyme
increased
significantly
in
the
100μg/L
treatment
at
five
weeks
compared
to
the
control,
while
no
significant
differences
were
recorded
in
the
25μg/L
treatment.
All
fish
within
the
25μg/L
were
male
(testes
all
containing
spermatozoa
and
spermatogenic
cells)
while
only
two
in
the
100μg/L
showed
both
pre-­‐vitellogenic
and
vitellogenic
oocytes
therefore
classified
as
female
compared
to
the
control
(6
female
and
4
male).
Only
the
100μg/L
treatment
caused
effects
to
the
fish.
Figure
4:
zebra
fish
A)
control
at
7
dpf,
B)
DBP
(100μg/L)
7
dpf
(Ortiz
–
Zarragoitia
et
al.,
2006)

27. 27
In
the
second
experiment
10
adult
female
zebra
fish
were
exposed
via
water
for
15
days
to
100
and
500μg/L
of
DBP.
After
15
days
of
exposure
each
female
was
paired
with
two
males
in
untreated
water
and
left
to
reproduce
for
two
to
three
days.
After
spawning
the
female
fish
were
sacrificed
and
liver,
brain
and
ovary
analysis.
Embryos
produced
during
spawning
were
gathered
and
placed
into
the
same
(treatment)
groups
as
their
female
parent
and
then
transferred
to
untreated
water
for
27
days.
The
number
of
eggs
produced
by
the
treated
females
did
not
differ
from
the
numbers
of
the
control.
However,
mortality
showed
a
significant
dose
dependent
relationship
such
as
in
the
highest
treatment
where
70%
mortality
was
recorded
after
25
days.
VTG
expression,
liver
VTG
protein
levels,
oocytes
and
ovary
development
showed
no
significant
difference
compared
to
the
control.
Both
experiments
incorporated
mortality
of
young
zerbra
fish,
however
exposure
to
phthalates
pre
fertilization
increased
the
mortality
where
as
exposure
post
hatch
showed
no
affect
on
mortality.
Aoki
et
al.,
(2011)
conducted
the
most
recent
in
vivo
study
on
DBP.
They
chose
adult
male
three-­‐spined
stickle
back
(Gasterosteus
aculetaus).
Fish
were
exposed
through
water
for
22
days
to
nominal
concentrations
of
50
and
100μg
DBP/L.
Throughout
the
experiment
the
concentrations
of
DBP
were
measured
every
three
to
four
days
(water
samples
ran
through
gas
chromatography
and
mass
spectroscopy)
where
it
was
found
that
the
actual
concentration
was
much
lower
than
their
original
calculated
input.
Mean
concentrations
of
15
and
35
μg/L
were
recorded
at
the
50
and
100μg/L
tanks,
respectively.
There
was
no
significant
difference
in
weight,
length
or
gonado-­‐somatic
index
for
either
treatment
group
compared
to
the
control.
They
did
find
that
testosterone
levels
and
oxidised
testosterone
levels
were
significantly
higher
in
the
35μg/L
treatment
group
(P<0.05)
compared
to
the
control.
Spiggin
(protein
glue)
was
also
measured
in
the
kidneys,
where
it
was
found
to
have
a
negative
correlation
with
DBP
concentration
with
only
the
highest
DBP
concentration
showing
a
significant
decrease
in
spiggin
(P<0.011).
A
slight
delay
in
nest
building
behavior
of
those
fish
in
the
35μg/L